Effect of a Constructed Wetland on Disinfection Byproducts: Removal

Environ. Sci. Technol. 2000, 34, 2703-2710

Effect of a Constructed Wetland on Disinfection Byproducts: Removal Processes and Production of Precursors COLLEEN E. ROSTAD,* BARBARA S. MARTIN, LARRY B. BARBER, AND JERRY A. LEENHEER U.S. Geological Survey, Building 95, Denver Federal Center, Denver, Colorado 80225 STEPHEN R. DANIEL Chemistry/Geochemistry Department, Colorado School of Mines, Golden, Colorado 80401

The fate of halogenated disinfection byproducts (DBPs) in treatment wetlands and the changes in the DBP formation potential as wastewater treatment plant (WWTP)-derived water moves through the wetlands were investigated. Wetland inlet and outlet samples were analyzed for total organic halide (TOX), trihalomethanes (THM), haloacetic acids (HAA), dissolved organic carbon (DOC), and UV absorbance. Removal of DBPs by the wetland ranged from 13 to 55% for TOX, from 78 to 97% for THM, and from 67 to 96% for HAA. The 24-h and 7-day nonpurgeable total organic halide (NPTOX), THM, and HAA formation potential yields were determined at the inlet and outlet of these wetlands. The effect of wetlands on the production of DBP precursors and their DBP-formation potential yield from wastewater was dramatic. The wetlands increased DBP yield up to a factor of almost 30. Specific changes in the DOC precursors were identified using 13C NMR spectroscopy.

Introduction As a result of widespread chlorine disinfection of drinking water, water-borne diseases such as typhoid, cholera, and dysentery no longer pose a threat to public health in the United States (1). Despite these direct benefits, chlorination of drinking water is not without risks. Recent studies have focused on increased incidence of spontaneous abortions or stillbirths (2, 3) or latent effects, such as cancer, from lifelong exposure to low concentrations of disinfection byproducts (DBPs) (4). Disinfection byproduct formation potential (DBP-FP) is dependent on factors such as the character, source, and concentration of the dissolved organic carbon (DOC) and disinfection reaction time, temperature, pH, and chlorine residual (5-7). In response to potential health hazards posed by DBPs, the Stage 1 Disinfection/Disinfection By-Product (D/DBP) Rule was promulgated (8). As a result, many municipal drinking water and wastewater utilities have explored methods to reduce concentrations of DBPs by utilizing mechanisms that reduce DBP precursors. In the arid southwestern United States, DBPs are of particular * Corresponding author e-mail: [email protected]; phone: (303)236-3971. 10.1021/es9900407 CCC: $19.00 Published on Web 05/31/2000

 2000 American Chemical Society

concern because sources of water are scarce, and many cities are evaluating wastewater reclamation and reuse to supplement their growing municipal, agricultural, and industrial water demands (9). Wastewater reclamation typically consists of treatment at a wastewater treatment plant followed by direct return to surface water or infiltration into groundwater that may subsequently become drinking water sources. Most discharge permits require disinfection, which can produce significant DBP concentrations. In recent years, there has been increasing interest in treatment wetland systems to accomplish the final polishing of wastewater treatment effluent (10-13). The fate of DBPs in treatment wetlands has not been studied extensively. Wetlands are highly productive and tend to increase concentrations and change characteristics of dissolved organic carbon (DOC) as the result of internal loading (15). Consequently, the capacity for wetland treatment systems to affect wastewater DBP-FP is an important question relating to human health, environmental hazards, and appropriate wetland treatment technology. The Tres Rios Demonstration Constructed Wetlands near Phoenix, AZ, is a well-documented wetland in terms of hydraulic retention time, hydraulic loading rate, operating depths, evapotranspiration losses, and infiltration losses (16). The site provides an opportunity to investigate, under carefully controlled conditions, the fate of DBP-FP in wetlands.

Site Description The Tres Rios Wetlands receive tertiary-treated sewage effluent from the nearby 91st Ave Wastewater Treatment Plant (WWTP), southwest of Phoenix, AZ. These wetlands consist of three distinct sites: the Hayfield site, the Cobble site, and the Research Cells (Figure 1). The 6-acre Hayfield site consists of two wetland cells of equal size and shape. Cell HS1 and HS2 differ in the number of deep zones, although the surface area of deep zone is the same. Deep zones are areas designed to have open water due to a water depth of up to 1.5 m (17). Between deep zones are vegetated marsh areas containing bulrush (Scirpus spp.) (17). WWTP effluent flows into each cell from the Hayfield site inlet box (HSin). Water flows through each cell to a separate outlet (H1out and H2out), which recombines to flow into a riparian zone (HScomb) discharging into the Salt River. The Cobble site, located within the Salt River flood plain, consists of two cells of equal size, which run in parallel. The northern cell, CS1, is not lined, and the bottom consists of native alluvial sand and gravel. The southern cell, CS2, is lined with an impervious clay-loam soil to a depth of 0.15 m (17). Each cell discharges to a outlet weir (CS1out or CS2out), which discharges into the Salt River channel. The Research Cells are 12 rectangular cells labeled R1-12. Areas of open water and vegetation and percent deep zone (11-35% open water) vary (17), and inflow hydraulic loading rate (HLR) and depth can be controlled. At the time of sampling, effluent from the WWTP was equally distributed to the 12 research cells, which operated in parallel (17).

Methods Sampling. Grab samples for total organic halide (TOX), trihalomethane (THM), and haloacetic acid (HAA) analyses were collected at wetland cell inlets and outlets between April 1997 and February 1998. A background sample was taken upstream from the site from the Salt River groundwater alluvium. Tap water from the WWTP and samples from the Salt and Gila Rivers at various locations also were collected. VOL. 34, NO. 13, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 1. Diagrams of the (a) Hayfield site, (b) Cobble site, and (c) Research Cells. The TOX samples were analyzed in duplicate in accordance with Method 5320 (6). Samples with greater than 30% relative difference between the duplicate measurements were reanalyzed. Uncertainty in each TOX determination was approximately 25%, and relative percent difference between duplicates averaged 24%. Analysis for THM, haloacetonitriles (HAN), haloketones (HK), and several other organic compounds was performed in accordance with U.S. Environmental Protection Agency (EPA) Method 551 (18). Uncertainty in each THM determination was approximately 25%, and relative percent difference between duplicates averaged 30%. Analysis for HAA was performed in accordance with EPA Method 552.1 (19). Uncertainty in each determination was approximately 30%, and relative percent difference between duplicates was less than 40%. Compound identifications were verified by GC/MS. 7-Day Ultimate DBP Formation Potential. Samples for DBP-FP analysis were collected in July 1997. 7-Day nonpurgeable total organic halide (NPTOX), THM, and HAA formation potential (FP) analyses were performed according to Method 5710 (6), with samples incubated at 25 °C at pH 7.0 with phosphate buffer. Samples also were collected for ultraviolet (UV) analysis. Specific UV absorbance (SUVA), the ratio of absorbance at 254 nm (UV254) per meter to DOC concentration, has been previously correlated with DBP precursors and DBP-FP (5). 2704

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24-h Simulated Distribution System DBP Formation Potential. A 24-h simulated distribution system formation potential (SDS-FP) analysis (6) was performed to simulate chlorination conditions at the City of Phoenix drinking water plant. Water is usually chlorinated at 2.0-2.5 mg/L free chlorine, residence time of water in the distribution system is about 24 h. and chlorine residual ranges from 0.3 to 0.9 mg/L (20). Our samples were adjusted to pH 7.5 and dosed (without buffer) with sodium hypochlorite to result in an initial free chlorine concentration of 2.25 mg/L, incubated at 25 °C, with a residual free chlorine of 0.1-1.0 mg/L after 24 h. Chlorination was quenched and samples analyzed as above. NPTOX samples were acidified, purged with nitrogen, and analyzed as above. Errors in the 24-h and 7-day measurements are approximately (25%, based on replicate analyses. Relative percent difference between duplicate samples was less than 40%. DOC was analyzed by heated persulfate oxidation using an Oceanographic International Carbon Analyzer. UV was measured at 254 nm in a 5-cm cell. Because formation potential is dependent upon DOC concentration, the NPTOXFP, THM-FP, and HAA-FP are reported in terms of formation potential “yield”, that is, micrograms of DBP formed per milligram of DOC precursor material in the water (µg of DBP/ mg of DOC).

Preparative Scale DOC Isolation. About 80 L were collected from wetland inlets and outlets for fulvic acid fractionation and characterization. Samples were fieldfiltered through 25- and 0.9-µm glass cartridge filters. Before isolation, DOC concentrations were determined. Fulvic acid was isolated using Duolite A-7 anion-exchange resin and Amberlite XAD-8 resin using procedures described by Leenheer and Noyes (21). 13C NMR Spectroscopy. Solid-state, cross-polarization, magic-angle-spinning (CP/MAS) 13C NMR spectra of fulvic acid isolates were obtained using a 200-MHz Chemagnetics CMX spectrometer with a 7.5-mm-diameter probe and a spinning rate of 5000 Hz. Acquisition parameters for the freeze-dried samples included a contact time of 1 ms, pulse delay of 1 s, and pulse width of 4.5 µs for the 90° pulse.

Results and Discussion Field Observations. Concentrations of TOX (Table 1) at Tres Rios ranged from about 75 to over 200 µg/L. TOX in the WWTP effluent is within range of other wastewaters previously reported: 85 µg/L in Washington, DC; 87 µg/L in Phoenix, AZ; 130 µg/L in Orange City, CA; 190 µg/L in Palo Alto, CA; and 170-290 µg/L in North Carolina (22, 23). TOX in the WWTP tertiary treatment effluent as it enters the wetland is about 200 µg/L, relatively high in comparison to other wastewater effluents. TOX in Phoenix tap water, about 160 µg/L, is relatively low as compared to the TOX range for finished drinking waters of 150-500 µg/L (24) with typical measurements in the range of 200-300 µg/L (23). During more intensive sampling, TOX concentrations decreased longitudinally along cell H1, while there was no significant change in concentration along a lateral profile (25), indicating fairly uniform flow through the wetlands. Distribution of THM (Table 1) was typical for chlorinated natural waters (26). For comparison, Phoenix tap water had nearly 70 µg/L total THM; whereas, the Gila River water had 0.1 µg/L total THM. Concentrations of HAN, HK, and other compounds were generally less than 5 µg/L in the wetland inlets and usually not detected at the outlets. Tap water concentrations of these compounds were generally greater than those found in the wetlands. Reported values of total THM concentrations range from 1.5-13 µg/L in secondary treated wastewater (22) to around 20 µg/L in tertiary treated wastewater (27). Total THM concentrations in finished drinking waters range from 1.4 to 97 µg/L (26). Concentrations determined in this study, for tertiary treated wastewater and tap water (68 and 11 µg/L), fall within these ranges. Stage I of the promulgated D/DBP rule will lower the maximum contaminant level (MCL) for total THM concentrations from 100 to 80 µg/L (8). Total HAA (Table 1) were consistently higher at the wetland inlets than at the outlets. In comparison, total HAA in Phoenix tap water was 75 µg/L and in the Gila River water was 0.4 µg/L. Total HAA were usually greater than total THM, total HAN, and total HK. Reported literature values for total HAA concentrations from WWTP effluents range from 2.1 to 63 µg/L (23, 26). Again, HAA concentrations at the Tres Rios wetland inlets are at the higher end of this range, which is attributed to the fact that this water is chlorinated and dechlorinated as the final step of tertiary treatment. Total HAA concentrations in chlorinated drinking waters averaged 77 µg/L in North Carolina (23) and ranged from 1 to 60 µg/L in Utah (26). The MCL for total HAA under the promulgated Stage I D/DBP rule is 60 µg/L (8). The TOX, THM, HAN, HK, and HAA all decreased as the wastewater moved through the various wetlands. Removal of TOX was 35% in HS1, 34% in HS2, 55% in CS1, and 43% in CS2. TOX removal from the research cells ranged from 13 to 52% and averaged 31%. A variation of the statistical t-test (28), employed to determine the difference between means

for two independent samples, found TOX at the wetland inlets and outlets to be significantly different at the 95% confidence level. Differences in removal between the Hayfield and Cobble sites is probably due to structural differences that allow more inlet water to infiltrate the ground in the Cobble site (16). Hydraulic loading of the cobble site is considerably higher to compensate for heavy losses due to infiltration. The average of about 40% removal of the TOX is similar to the 40-50% TOX removal reported during soil aquifer treatment of tertiary treated wastewater effluent (27). Compounds comprising TOX include many large, polar, more complex compounds (29), but the majority of TOX loss was attributed to loss of THM, HAN, HK, and HAA. The remaining portion of the TOX is mostly nonvolatile (29). These compounds are more likely to sorb onto sediments and organic matter in soil, given their high partition coefficients (30). Wetland sediments could effectively sequester hydrophobic TOX because they tend to have a high organic carbon content and are constantly supplied with organic carbon from vegetative die-off (10). Microbial degradation and photolysis also may remove TOX (31). At the Tres Rios site, nearly 60% of the inlet TOX was present at the wetland outlet. In the case of the Cobble site, TOX also may be removed by groundwater infiltration. Reduction of THM in the wetland ranged from 83% in HS1, 84% in HS2, 87% in CS2, to 97% in CS1. Removal from the Research Cells ranged from 78 to 93% and averaged 84%. Volatilization was considered the most likely route of THM removal due to high vapor pressures at ambient temperatures, which facilitate evaporation from an open water body (31). Chloroform has a relatively high Henry’s law constant (31) and evaporates during wastewater treatment in less than 2 h (32). Average THM reduction was 87% for all of the wetland cells, which is consistent with THM removal reported from water during soil aquifer treatment (27). Other mechanisms for THM removal include biodegradation, photolytic degradation, and abiotic hydrolysis (31). For most THM, aerobic biodegradation half-life is in the range of weeks, while volatilization half-life is in the range of hours to days (31). Measured hydraulic retention times (HRTs) for the wetland cells range from 1.1 to 12 days (16). At the time of sampling, HRTs for HS1 and HS2 were 3.4 and 3.3 days, while HRTs for CS1 and CS2 were 2.1 and 2.0 days (16). Since hydraulic retention time in the wetland cells is short, volatilization likely dominates over biodegradation as a removal mechanism. Degradation through abiotic photolysis and hydrolysis may occur, but rates are very slow. The hydrolysis half-life of bromodichloromethane at 25 °C is approximately 137 yr (31). Removal of HAA was comparable to that of THM: 85% in HS1, 89% in HS2, 92% in CS1, and 79% in CS2 and the Research Cells ranged from 67 to 96% and averaged 84%. The water-soluble HAA are unlikely to be removed by adsorption to sediments and are not as likely to volatilize based on their low vapor pressures and Henry’s law constants (32). Abiotic hydrolysis also is unlikely (31). A possible mechanism for the removal of HAA from water is microbial degradation. Chloroacetic acid was found to be degraded with a half-life of 3.4 h by Methylosinus trichosporium OB-3b, a soil methylotroph (33). HAA also have been degraded by haloacid dehalogenase, a bacterial enzyme, into hydroxyalkanoic or oxoalkanoic acids (34). Other bacteria such as Xanthobacter antotrophicus degrade HAA under aerobic conditions, with half-lives of 1-2 days (34). Photooxidation of HAA may occur with simultaneous photoreduction of iron oxyhydroxide minerals, at slow rates, into haloacids and CO2 (35). Research is needed to determine the fate of HAA in environmental systems, especially since (as DBPs) their concentrations may exceed THM in tertiary treated wastewater. VOL. 34, NO. 13, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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TABLE 1. Concentration of Selected Disinfection Byproducts (in µg/L) at the Tres Rios Sampling Sites

site Tap CSin CS1out CS2out HSin HS1out HS2out HScomb R1in R1out R2out R3out R4out R5out R6out R7out R8out R9out R10out R11out R12out 91st Ave 115th Ave Bullard Ave Gila River Groundwater Blank a

bromo1,1,11,11,2tritetra- 1,1,1bromo- dibromochloro- dichloro- trichloro- bromo- dibromodichloro- chloro- dibromo- trichloro1,2trichloro- dichloro- dibromo- chloro- chlorotrichloro- dichloro- chloro- bromo- total acetic acetic acetic acetic acetic total aceto- aceto- acetonitro- dibromo- 2-propan- 2-propan- 3-chloro- ethyl- ethyl- chloroTOX form methane methane form THM acid acid acid acid acid HAA nitrile nitrile nitrile methane ethane one one propane ene ene ethane 156 170 76 97 214 140 101 127 178 130 113 155 127 120 140 128 142 128 111 85 102 190

23 8.1 0.3 1.4 8.2 1.8 1.8 2.7 7.9 1.3 1.9 2.1 1.7 1.4 1.9 1.4 1.0 1.8 1.7 1.8 0.8 5.4

23 2.3 nd nd 1.5 nd nd nd 1.0 nd 0.2 0.2 0.1 0.2 0.1 0.1 nd 0.2 0.1 nd nd 1.1

19 0.4 nd nd 0.7 nd nd nd 0.7 nd nd nd nd nd nd nd nd nd 0.1 nd nd 0.4

2.6 0.1 nd nd 0.5 nd nd nd 0.8 nd nd nd nd nd nd nd nd nd nd nd nd nd

68 11 0.3 1.4 11 1.8 1.8 2.7 11 1.3 2.1 2.4 1.8 1.6 2.0 1.5 1.0 1.0 1.8 1.8 0.8 6.8

2.9 2.6 nd nd 4.3 nd nd nd 40 nd nd nd nd nd nd nd nd nd nd nd nd 19

50 17 2.2 2.2 18 2.3 2.4 7.8 27 10 5.8 14 21 7.1 3.6 3.6 16 23 2.8 3.0 16 29

9.8 3.7 0.5 0.4 2.3 0.3 0.2 0.5 3.0 nd nd 1.2 1.3 0.1 0.5 1.3 0.2 0.3 nd nd 1.0 9.9

12.0 7.9 4.5 5.6 5.8 6.1 8.1 nd nd nd nd nd nd nd nd nd nd nd nd nd nd nd

nda 0.2 nd nd 0.4 nd nd nd nd nd nd nd nd nd nd nd nd nd nd nd nd nd

75 31 7.2 8.2 31 8.7 11 8.3 70 10 5.8 15 22 7.2 4.1 4.9 16 23 2.8 3.0 17 57

4.0 2.6 nd nd 3.4 nd nd nd 3.0 nd nd nd nd nd nd nd nd nd nd nd nd 0.6


1.4 nd nd nd 0.2 nd nd nd 0.1 nd 0.2 nd 0.2 nd nd nd nd 0.2 nd nd nd nd


11 0.4 nd nd 1.5 nd nd nd 1.0 nd nd nd nd nd nd nd nd nd nd nd nd nd



0.1 nd nd nd nd 0.3 0.1 0.1 nd 0.1 0.1 nd 0.1 0.1 nd nd 0.1 0.2 0.1 nd nd 0.1

nd 0.6 nd nd 0.5 nd nd nd 0.5 nd nd nd nd nd nd nd nd nd nd nd nd 0.6

0.1 0.1 nd 0.2 0.7 0.1 nd nd 0.6 0.1 0.2 0.2 0.1 0.1 nd 0.2 nd 0.1 nd 0.1 nd 0.2


108

3.8

0.5

0.3

nd

4.7

nd

12

7.0

nd

nd

19

0.3

nd

nd

nd

nd

0.1

0.4

nd

0.3

0.2

nd

98

2.3

0.3

0.1

nd

2.7

46

7.9

3.5

nd

nd

58

nd

nd

nd

nd

nd

nd

nd

nd

nd

nd

nd

50

0.1

nd

nd

nd

0.1

nd

0.1

0.2

nd

nd

0.4

nd

nd

nd

nd

nd

nd

nd

nd

nd

nd

nd

91

0.2

nd

nd

nd

0.2

82

14

nd

nd

nd

96

nd

nd

nd

nd

nd

nd

nd

nd

nd

nd

nd

9

nd

nd

nd

nd

nd

nd

1.6

0.1

2.7

nd

4.4

nd

nd

nd

nd

nd

nd

nd

nd

nd

nd

nd

nd, not detected.

TABLE 2. Summary of 24-h and 7-day NPTOX, THM, and HAA Formation Potential Yield, SUVA, and DOC from Selected Wetland Sites 24-h

HSin HS1out HS2out CSin CS1out CS2out a

SUVA

DOC

µg NPTOX/ mg DOC

µg THM/ mg DOC

µg HAA/ mg DOC

µg NPTOX/ mg DOC

µg THM/ mg DOC

7-day µg HAA/ mg DOC

abs (L)/ m (mg)

mg/L

16 20 26 2.0 58 22

0.3 4.1 5.2 0.0 1.9 3.3

1.1 1.7 1.4 0.0 1.6 0.9

98 210 170 26 170 220

75 77 78 54 62 94

7.8 12 14 8.4 9.9 37

1.7 2.9 2.7 2.0 2.2 3.3

7.65 7.35 7.0 7.65 9.93 9.93

abs, absorbance.

The Tres Rios wetlands are removing much of the inlet DBPs. Reduction of various DBPs from 30 to 97% illustrates that, in addition to providing riparian habitat, wetlands may improve water quality. In terms of water reuse, passive wetland treatment of WWTP effluent may be an alternative to surface spreading basins or direct groundwater infiltration. DBP Formation Potential Yields. Results for NPTOX-FP, THM-FP, and HAA-FP are listed in Table 2. Wastewater is chlorinated and dechlorinated prior to release into the wetlands, lowering its DOC reactivity, so yields from the inlet wastewater were expected to be low. The low FP at the CSin may be due to longer transit time between the plant and the wetland during which time any residual free chlorine would have exhausted most DOC reactivity. No free chlorine was detected in any of the FP samples, although there were trace levels in the inlet samples. In reclaimed water, also chlorinated and dechlorinated prior to release, 7-day TOX yields ranged from 13 to 50 µg of TOX/mg of DOC (22). Wetland outlet 7-day NPTOX-FP yields varied over a rather narrow range when compared to other studies. Typical TOX yields from FP studies on surface waters ranged from about 80 to 500 µg TOX/mg DOC, and yields of NPTOX (which excludes THM) ranged from 170 to 350 µg NPTOX/mg DOC (36). TOX yields of raw drinking waters from seven U.S. cities ranged from 170 to 300 µg TOX/mg DOC (24). TOX yields from five surface waters ranged from 140 to 230 (average 190) µg TOX/mg DOC for fulvic acids and ranged from 230 to 290 (average 260) µg TOX/mg DOC for humic acids (24). Differences in yield are usually attributed to differences in the character or source of the DOC. The THM-FP yields at the wetland outlets were much higher than at the inlets. Other 7-day THM-FP analyses have yielded results ranging from 32 to 190 µg THM/mg DOC. Yields of THM from rivers ranged from 47 to 190 µg THM/mg DOC (36), from raw drinking waters ranged from 41 to 68 µg THM/mg DOC (24), and from groundwaters ranged from 25 to 83 µg THM/mg TOC (37). Yields of the wetland 7-day HAA-FP were low and usually much lower than the NPTOX-FP and THM-FP. Other reported HAA-FP yields ranged from 14 to 22 µg HAA/mg TOC (38). Differences in HAA-FP between wetland inlets and outlets were small, except at CS2out, which also had the highest NPTOX-FP and THM-FP. Although DOC usually increased slightly as water moved through the wetland (15), DBP-FP yield increased even more. Wetland outlets produced more DBP upon chlorination than inlets in both the 24-h and 7-day analysis. Note that in all cases, the 7-day FP was greater than the 24-h FP because increasing the reaction time increases DBP yield (39). The wetlands are contributing DOC to the wastewater (15), and that DOC is highly reactive in producing DBPs. The 24-h yield of both NPTOX-FP and HAA-FP at the Hayfield site increased by factors of 1.3-1.6 between the inlet and the outlet. However, the 24-h THM-FP yield

increased by a factor of 14-17. This indicates that the DOC contributed by the Hayfield wetland is more reactive, i.e., it produces more THM and more quickly than DOC in wastewater. Largest increases in FP were observed in the 24-h NPTOX at the Cobble site, which increased by a factor of 11 at CS2 and 29 at CS1, between the wetland inlets and outlets. The 7-day Hayfield site NPTOX-FP and HAA-FP increases were similar to the 24-h increases. But the 7-day THM-FPs were quite different than the 24-h THM-FPs. Increases in 7-day THM-FPs more closely followed increases in the HAA-FPs. THM-FPs increased by a factor of 1-1.7, while the HAA-FP increased by a factor of 1.2-4.4. NPTOXFP increased by a factor of about 2 at the Hayfield sites and by a factor of about 7.5 at the Cobble sites. As discussed above, losses to infiltration at the Cobble site, especially CS1, are great, resulting in a longer hydraulic retention time than at the Hayfield site. In general, DOC contributed by the wetland is very reactive in the production of DBPs. Another important comparison is the relationship between the 24-h and 7-day analyses (Figure 2), although differences in their initial conditions should be considered. Note that HAA are shown separately, although they constitute part of the NPTOX. In both the 24-h and 7-day samples, NPTOX was the most abundant, and HAA were at a consistent minimal percentage (4-7%). However, the proportion of THM formed changed over time and through the wetland. This indicates that reactions of DOC with chlorine to produce NPTOX and HAA are initially faster than those producing THM. A study by Rathbun (39) on NPTOX and THM production over time indicated similar results, where it was observed that 50% of the final 7-day NPTOX concentration was formed in 2.7 h, while it took 16.2 h to produce 50% of the final 7-day THM concentration. The proportion of HAA formed over time with respect to NPTOX was nearly constant in both 24-h and 7-day samples, averaging 5.5% of the NPTOX, excluding the 7-day CSin and CS2out outliers. This implies that HAA form at a constant rate over the 7-day period, similar to the formation rate of NPTOX. In contrast, both 24-h and 7-day THM yields varied widely with respect to NPTOX. Formation Potential and Specific UV Absorbance (SUVA). Absorption at 254 nm is indicative of aromatic carbon and phenolic precursor compounds from which DBPs form (40). Therefore, SUVA may correlate with DBP yield (5, 7, 41). Most of the wetland inlets had lower SUVA (Table 2) than the outlets and produced less NPTOX, THM, and HAA over a 24-h and 7-day period. This may be due to changes in DOC character or reactivity between the wetland inlet, with essentially exhausted reactivity, and the wetland outlet, with autochthonous, highly reactive DOC. SUVA was compared with DBP-FP yields to examine possible correlations between FP yield and aromaticity. Linear correlations between the 7-day DBP-FPs and SUVA indicate that aromaticity correlates with NPTOX-FP (r 2 ) 0.67), THM-FP (r 2 ) 0.56), and VOL. 34, NO. 13, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 2. Comparison of NPTOX-FP, THM-FP, and HAA-FP yields between 24-h and 7-day samples in samples from HSin and HS1out. HAA-FP (r 2 ) 0.74, excluding CS2out). No correlation was found between 24-h NPTOX-FP (r 2 ) 0.005) or HAA-FP(r 2 ) 0.08) and SUVA, although SUVA correlated with 24-h THMFP (r 2 ) 0.60). This indicates that aromatic carbon may not be solely responsible for NPTOX and HAA production in the initial 24 h of reaction but may produce DBPs later in the reaction period. Although SUVA is used as a predictor of THM-FP, it does not appear applicable for HAA-FP, especially in the 24-h time frame typical of drinking water plants. The wetland is both a source and a sink for DOC. Some of the influent DOC is likely removed from the water by processes such as sorption and microbial degradation. However, DOC is also entering the water through vegetative die-off. The different sinks and autochthonous sources of DOC in the wetland may be responsible for changes in reactivity of DOC between the inlets and the outlets. Although a major reaction believed to be responsible for production of THM is the reaction of free chlorine with m-dihydroxybenzene structures (40), other compound classes also may be responsible, including ketones, β-keto acids, and tannins. Of these reactants, only the dihydroxybenzene and tannin structures will absorb at 254 nm (42) and contribute to SUVA. Because the dominant absorbances for ketones and β-keto acids do not maximize near 254 nm, no strong correlations can be made between SUVA and THM formation through these particular reaction pathways. 13C NMR Characterization of Fulvic Acid. The fulvic acid fraction that constitutes about 40% of the DOC for both wetland inlet and outlet was isolated and characterized by 13C NMR spectrometry to obtain additional insights into the organic structural groups that might be reactive with chlorine 2708

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FIGURE 3. 13C NMR spectra for fulvic acid isolates from the Hayfield site: (A) wetland inlet and (B) wetland outlet. and affect UV absorbance. The 13C NMR spectra for fulvic acid isolates from the wetland inlet and outlet are shown in Figure 3.

Both inlet and outlet spectra have significant aromatic carbon content as evidenced by peaks between 100 and 160 ppm. These aromatic carbons in treated sewage effluents are primarily sulfonated anionic surfactant metabolites (43). The peak near 140 ppm is indicative of the aromatic carbon to which the sulfonic acid group is attached. The critical difference between these two spectra is the addition of aromatic carbon in the 100-120 ppm and 140-160 ppm regions. This type of aromatic carbon is indicative of condensed tannins (44) that are highly reactive with chlorine (45). Tannins are rich in m-dihydroxybenzene structures because of the phloroglucinol ring structure found in condensed tannins. Both tannins and their humic substance degradation products are highly colored because of multiple aromatic ring systems that are in conjugation, and it is this property that explains the large increase in SUVA and DBPFP with a relatively small increase in aromaticity and DOC concentration. Generally, treatment by the Tres Rios wetlands increased the DBP formation potential yield of the influent wastewater. The 7-day and 24-h formation potentials of all DBPs were increased by passage through the wetland by factors from 1 to almost 30. Results for NPTOX and THM yields from this investigation are consistent with previous studies. Although the Tres Rios wetlands substantially remove DBPs (25), the wetlands increase DBP formation potential relative to the treated sewage effluent. The 7-day NPTOX-FP, THM-FP, and HAA-FP yields were found to correlate linearly with SUVA, while the 24-h NPTOX-FP and HAA-FP did not. This indicates structural differences in the portion of the DOC producing DBPs after 24 h and 7 days. Finally, specific portions of the DOC responsible for these FP increases were documented with 13C NMR spectra. The DBP-reactive natural organic matter added by the constructed wetlands may well be removed by iron and alum flocculation during drinking water treatment.

Acknowledgments This research was supported by the U.S. Bureau of Reclamation. We acknowledge Ted Noyes from U.S. Geological Survey; Eric Stiles from U.S. Bureau of Reclamation; and Roland Wass, Wes Camfield, and Ronnie Elkins from City of Phoenix. Use of trade names is for identification purposes only and does not constitute endorsement by the U.S. Geological Survey.

Literature Cited (1) Johnson, J. D.; Jolley, R. L. In Water Chlorination: Chemistry, Environmental Impact and Health Effects, Vol. 6; Jolley, R. L., Condie, L. W., Johnson, J. D., Katz, S., Minear, R. A., Mattice, J. S., Jacobs, V. A., Eds.; Lewis Publishers: Chelsea, MI, 1990; pp 21-27. (2) Swan, S. H.; Waller, K.; Hopkins, B.; Windham, G.; Fenster, L.; Schaefer, C.; Neutra, R. R. Epidemiology 1998, 9, 126. (3) Dodds, L.; King, W.; Woolcott, C.; Pole, J. Epidemiology 1999, 10, 233. (4) Zierler, S.; Feingold, L.; Danley, R. A.; Craun, G. F. In Water Chlorination: Chemistry, Environmental Impact and Health Effects, Vol. 6; Jolley, R. L., Condie, L. W., Johnson, J. D., Katz, S., Minear, R. A., Mattice, J. S., Jacobs, V. A., Eds.; Lewis Publishers: Chelsea, MI, 1990; pp 399-410. (5) Minear, R. A.; Amy, G. L. In Water Disinfection and Natural Organic Matter: Characterization and Control; Minear, R. A., Amy G. L., Eds.; American Chemical Society, Washington, DC, 1996; pp 1-9. (6) Greenberg, A. E., Clesceri, L. S., Eaton, A. D., Eds. Standard Methods for the Examination of Water and Wastewater, 19th ed.; American Public Health Association: Washington, DC, 1995. (7) Najm, I. N.; Patania, N. L.; Jacangelo, J. G.; Krasner, S. W. J. Am. Water Works Assoc. 1994, 86, 98. (8) U.S. Environmental Protection Agency. Fed. Regist. 1998, 63 (241), 69390.

(9) Hammer, D. A. Constructed Wetlands for Wastewater Treatment: Municipal, Industrial, and Agricultural; Lewis Publishers: Chelsea, MI, 1989; 856 pp. (10) Kadlec, R. H.; Knight, R. L. Treatment Wetlands; Lewis Publishers: Boca Raton, FL, 1996; 848 pp. (11) U.S. Environmental Protection Agency. Constructed Wetlands for Wastewater Treatment and Wildlife Habitat; EPA 831-R93-005; U.S. Government Publication Office: Washington, DC, 1993; 174 pp. (12) Reed, S. C.; Crites, R. W.; Middlebrooks, E. J. Natural Systems for Waste Management and Treatment, 2nd ed.; McGraw-Hill: New York, 1995. (13) McKinlay, R. G.; Kasperek, K. Water Res. 1999, 33, 505-511. (14) Pinney, M. L.; Westerhoff, P. K.; Baker, L. Water Res. 2000, 34, 1897-1911. (15) Barber, L. B.; Leenheer, J. A.; Noyes, T. I.; Stiles, E. A. Transformation of Dissolved Organic Matter in Constructed Wetland Systems; U.S. Bureau of Reclamation Project Report, Wetlands and Riparian Habitat Initiative; U.S. Government Printing Office: Washington, DC, January 1999. (16) Wass, R. Tres Rios Demonstration Constructed Wetland Project 1996/1997 Operation and Water Quality Report; Tulleson, AZ, 1997; 250 pp. (17) CH2MHILL. Tres Rios Wetlands Research Plan; Report SWW70423.02.WK; Phoenix, AZ, 1995. (18) U.S. Environmental Protection Agency. Method 551. In Methods for the Determination of Organic Compounds in Drinking Water, Supplement 1, July 1990; PB91-146027, EPA-600/4-90/020; U.S. Government Printing Office: Washington, DC, 1990; pp 169200. (19) U.S. Environmental Protection Agency. Method 552.1. In Methods for the Determination of Organic Compounds in Drinking Water, Supplement 2, August 1992; PB92-207703, EPA600/R-92/129; U.S. Government Printing Office: Washington, DC, 1992; pp 143-172. (20) Schebbler, S., Superintendent, City of Phoenix Drinking Water Treatment Plant, personal communication, 1997. (21) Leenheer, J. A.; Noyes, T. I. U.S. Geol. Surv. Water-Supply Pap. 1984, No. 2230, 16 pp. (22) Barber, L. B., II; Brown, G. K.; Kennedy, K. R.; Leenheer, J. A.; Noyes, T. I.; Rostad, C. E.; Thorn, K. A. In Proceedings of the American Water Resources Association Symposium, Conjunctive Use of Water Resources, Aquifer Storage and Recovery, October 19-23, 1997; Kendall, D. R., Ed.; American Water Resources Association: Long Beach, CA, 1997; pp 261-272. (23) Singer, P. C.; Obolensky, A.; Grenier, A. J. Am. Water Works Assoc. 1995, 87, 83. (24) Reckhow, D. A.; Singer, P. C. J. Am. Water Works Assoc. 1990, 82, 173. (25) Martin, B. S. Effect of a Constructed Wetland on Disinfection By-Products and their Formation Potentials. M.Sc. Thesis, Colorado School of Mines, 1998, 109 pp. (26) Nieminski, E. C.; Chaudhuri, S.; Lamoreaux, T. J. Am. Water Works Assoc. 1993, 85, 98. (27) Amy, G.; Wilson, L. G.; Conroy, A.; Chahbandour, J.; Zhai, W.; Siddiqui, M. Water Environ. Res. 1993, 65, 726. (28) Ott, L. An Introduction to Statistical Methods and Data Analysis, 3rd ed.; PWS-Kent Publishing: Boston, MA, 1988; 945 pp. (29) Richardson, S. D. In The Encyclopedia of Environmental Analysis & Remediation, Vol. 3; Meyers, R. A., Ed.; John Wiley & Sons: New York, 1998; pp 1398-1421. (30) Mackay, D.; Shiu, W. Y.; Ma, K. C. Illustrated Handbook of Physical-Chemical Properties and Environmental Fate for Organic Chemicals, Vol. 1; Lewis Publishers: Boca Raton, FL, 1992; 697 pp. (31) Howard, P. H., Ed. Handbook of Fate and Exposure Data for Organic Chemicals, Vol. II: Solvents; Lewis Publishers: Chelsea, MI, 1990; 546 pp. (32) Verschueren, K. Handbook of Environmental Data on Organic Chemicals, 2nd ed.; Van Nostrand Reinhold: New York, 1983; 1310 pp. (33) Castro, C. E.; O’Shea, S. K.; Wang, W.; Bartnicki, E. W. Environ. Sci. Technol. 1996, 30, 1180. (34) Hashimoto, S.; Azuma, T.; Otsuki, A. Environ. Toxicol. Chem. 1998, 17, 798. (35) Pehkonen, S. O.; Siefert, R. L.; Hoffman, M. R. Environ. Sci. Technol. 1995, 29, 1215. (36) Rathbun, R. E. Arch. Environ. Contam. Toxicol. 1996, 31, 420. (37) Miller, R. E.; Randtke, S. J.; Hathaway, L. R.; Denne, J. E. J. Am. Water Works Assoc. 1990, 82, 49. VOL. 34, NO. 13, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

2709

(38) Summers, R. S.; Hooper, S. M.; Shukairy, H. M.; Solarik, G.; Owen, D. J. Am. Water Works Assoc. 1996, 88, 80. (39) Rathbun, R. E. Chemosphere 1997, 34, 2699. (40) Larson, R. A.; Weber, E. J. Reaction Mechanisms in Environmental Organic Chemistry; Lewis Publishers: Boca Raton, FL, 1994; 433 pp. (41) Korshin, G. V.; Li, C.; Benjamin, M. M. In Water Disinfection and Natural Organic Matter: Characterization and Control; Minear, R. A., Amy, G. L., Eds.; American Chemical Society: Washington, DC, 1996; pp 182-195. (42) Silverstein, R. M.; Bassler, G. C.; Morrill, T. C. Spectrometric Identification of Organic Compounds, 5th ed.; John Wiley & Sons: New York, 1991; 419 pp.

2710

9


(43) Field, J. A.; Leenheer, J. A.; Thorn, K. A.; Barber, L. B., II.; Rostad, C. E.; Macalady, D. L.; Daniel, S. R. J. Contam. Hydrol. 1992, 9, 55. (44) Porter, L. J.; Newman, R. H.; Foo, L. Y.; Wong, H. J. Chem. Soc., Perkin Trans. 1 1982, 1217. (45) Fam, S.; Stenstrom, M. K. J. Water Pollut. Control Fed. 1987, 59, 11.

Received for review January 13, 1999. Revised manuscript received April 3, 2000. Accepted April 12, 2000. ES9900407

Effect of a Constructed Wetland on Disinfection Byproducts: Removal

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